Advanced electrocatalytic redox processes for environmental remediation of halogenated organic water pollutants

Halogenated organic compounds are widespread, and decades of heavy use have resulted in global bioaccumulation and contamination of the environment, including water sources. Here, we introduce the most common halogenated organic water pollutants, their classification by type of halogen (fluorine, chlorine, or bromine), important policies and regulations, main applications, and environmental and human health risks. Remediation techniques are outlined with particular emphasis on carbon–halogen bond strengths. Aqueous advanced redox processes are discussed, highlighting mechanistic details, including electrochemical oxidations and reductions of the water–oxygen system, and thermodynamic potentials, protonation states, and lifetimes of radicals and reactive oxygen species in aqueous electrolytes at different pH conditions. The state of the art of aqueous advanced redox processes for brominated, chlorinated, and fluorinated organic compounds is presented, along with reported mechanisms for aqueous destruction of select PFAS (per-and polyfluoroalkyl substances). Future research directions for aqueous electrocatalytic destruction of organohalogens are identified, emphasizing the crucial need for developing a quantitative mechanistic understanding of degradation pathways, the improvement of analytical detection methods for organohalogens and transient species during advanced redox processes, and the development of new catalysts and processes that are globally scalable.

This journal is © The Royal Society of Chemistry 2023

Introduction
The industrial revolution led to unprecedented technological progress, human health, prosperity, wellbeing, and population growth, but it also caused climate change 1 and environmental pollution on a global scale.Technological innovations in consumer goods, and pharmaceutical, agricultural, and industrial applications released large quantities of halogenated organic pollutants into the environment, including Earth's waterways.9][20][21][22][23][24][25][26][27] They are organic compounds that contain the halogen atoms fluorine, chlorine, or bromine, possess solubility in natural water, and are harmful to the environment.Halogenated organic compounds, also called organohalogens, are typically classified into different categories depending on their structure or chemical properties.Halogenated organic water pollutants include aliphatic or aromatic halogenated hydrocarbons. 28Common organohalogens are shown in Fig. 1.Organohalogens serve as solvents, coatings, degreasing agents, biocides, medical propellants, plasticizers, hydraulic and heat transfer fluids, chemical synthesis intermediates, refrigerants, coolants, and flame retardants. 22,23,25,27rganohalogens possess exceptionally strong carbon-halogen bonds, leading to high heat resistance, low surface tension, high lipophilicity, and chemical inertness.Most organohalogens possess amphiphilic (ionic and neutral) properties and are xenobiotic, although natural organohalogens are known. 29ecause of their thermodynamically strong covalent C-X (X = F, Cl, or Br) bonds, these compounds were initially considered nonmetabolizable and nontoxic, 30 which turned out to be false, creating global human health risks and an urgent need for environmental remediation, particularly from water sources.
The discovery and manufacturing of organohalogens has led to revolutionary materials with high utility, such as non-stick Teflon cookware coatings, Rain-X water repellants, fire-retardant, water-proofing, and grease-resisting additives to upholstery and clothing, firefighting foams, and cleaning agents for electronics and microelectronics manufacturing.The widespread use of these chemicals has resulted in inadvertent or purposeful discharge into the environment. 31The high strength of carbon-halogen bonds inhibits biodegradation processes in nature, and leads to extended lifetimes and accumulation of chlorinated, brominated, and fluorinated organic compounds in the environment, animals and humans. 20olicy regulations and restrictions regarding the manufacturing and use of halogenated organics have been instated on the national level and worldwide.For example, restrictions on the production of polychlorinated biphenyls (PCBs) began in 1970s, followed by international implementations through the Stockholm Convention in 2004, which banned the production of PCBs, aiming to phase out PCBs in use by 2025, and ensuring environmentally sound management by 2028. 32Limitations on brominated compounds started in 2009 when the Stockholm Convention listed polybrominated diphenyl ethers as new persistent organic pollutants and banned their production and use. 33Currently, five specific groups of brominated flame retardants are listed in the Stockholm Convention: hexabromobiphenyl, hexabromocyclododecane (HBCD), and the commercial polybrominated diphenyl ethers octabromodiphenyl ether (octaBDE), pentabromodiphenyl ether (pentaBDE), and decabromodiphenyl ether (decaBDE). 33dditional policy efforts have been made to discontinue or limit the production and use of poly-fluorinated compounds.Significant New Use Rules were established in the U.S. to restrict the production and use of per-and polyfluoroalkyl substances (PFAS).The U.S. Environmental Protection Agency worked with leading chemical companies on a global phaseout of perfluorooctanoic acid (PFOA) through the 2010/2015 PFOA Stewardship Program to reduce emissions and residual content of PFAS. 34By 2009, perfluorooctane sulfonic acid (PFOS) and related compounds were listed under Annex B of the Stockholm Convention on Persistent Organic Pollutants. 35The U.S.
Environmental Protection Agency's notification levels for PFOA and PFOS are 5.1 and 6.5 ppt, respectively, and the response levels for PFOA and PFOS are 10 and 40 ppt, respectively. 36,37hlorofluorocarbons (CFCs) and hydrochlorofluorocarbons (HCFCs) are fully or partially halogenated hydrocarbons that are produced from methane, ethane, and propane.CFCs and HCFCs are nontoxic, non-flammable, and long-lived synthetic compounds that contain carbon, hydrogen, chlorine, and fluorine. 38Any production of CFCs was banned in 2010 through the Montreal Protocol to reduce CFC emissions into the atmosphere, where CFCs deplete stratospheric ozone. 39][42] Resistance to biodegradation causes long lifetimes and accumulation of organohalogens in the environment, together with global distribution via waterways. 43Bioaccumulation through food chains and direct uptake introduce organohalogens into the human body, where they have been linked to Parkinson's disease, 44 harm in cognitive function and development, reproductive, hormonal, and metabolic processes, and increased risk for cancer. 30,45,46Specific applications and associated risks of chlorinated, brominated, and fluorinated organic compounds are detailed below.
Chlorinated and brominated organic water pollutants.Chlorinated and brominated organic water pollutants are harmful to human health and the environment. 46,47They accumulate in living organisms. 46Chlorinated organic compounds found uses in metal working fluids, lubricants, flame retardants, and plasticizers. 48Polybrominated diphenyl ethers used in consumer, commercial, and industrial products, 75 from aerospace to food production. 76Fluoropolymer coatings can repel water and oil, and resist thermal, chemical and biological decomposition.Decades of heavy use have resulted in bioaccumulation and contamination of water, soil, animals, and people all over the world. 77PFAS accumulation has even been observed in Antarctica. 78For example, PFOA has a half life of 3.8 years in humans, and PFOS accumulates in the human liver and is trapped by bile acid transporters, resulting in a half life of 7.1 years in humans. 79PFAS exposure is harmful to human health. 80,81The extreme chemical stability of PFAS arises from numerous C-F bonds, which is beneficial for the durability of products, but problematic for the environment and ultimately human health, as PFAS resist biodegradation.
PFAS have been identified in ground water, surface water, drinking water, soil, indoor air, dietary sources, manufacturing locations, and wastewater treatment plants, as well as in human blood, serum, milk, urine, and placenta tissue. 82For example, the PFAS chemical PFOS bioaccumulates in fish that humans consume, making such fish a source of human PFAS exposure. 83Human PFAS exposure can additionally occur via dermal contact because PFAS can be taken up from oil-and water-resistant non-stick coatings on cookware and paper food packaging materials. 83Inhalation of indoor air and household dust is another route for human PFAS uptake. 83uman exposure to PFAS causes reproductive and immune system harm, increased risk of non-Hodgkin's lymphoma and kidney, testicular, prostate, breast, liver, and ovarian cancers, decreased vaccine response, increased risk of asthma in adolescents, increased risk of diabetes and hypertension in women, developmental delays in children, changes in liver enzymes, endocrine disruption, and increased cholesterol levels and/or risk of obesity; 80,81,[84][85][86][87][88][89] in unborn children, delayed mammary gland development, reduced response to vaccines, and lower birth weight have been observed (Fig. 2). 90,91PFAS also pose ecological risks through bioaccumulation, food chains, and toxicity in terrestrial and aquatic wildlife. 92

Remediation techniques for halogenated organic water pollutants
Maximum contaminant limits in water in the U.S. are set by the U.S. Environmental Protection Agency and published in the National Primary Drinking Water Regulations. 93For most chlorinated, brominated, and fluorinated organic pollutants, the allowable contamination threshold is on the order of parts per billion (ppb) or parts per trillion (ppt), depending on the compound and associated risks. 52,94To achieve safe levels, water remediation techniques are needed for effective removal and destruction of pollutants.Halogenated organic pollutants have a particularly high resistance against degradation because of their strong carbon-halogen bonds.As a result, only a few destruction techniques for complete mineralization of halogenated organic water pollutants have been commercialized, and the field is a very active research area. 27,95,96Several separation techniques have found commercial use to remove halogenated pollutants from water sources; separation only concentrates pollutants, 27 after which disposal by destruction is necessary.

Concentration vs. destruction techniques
Organohalogen concentrations differ vastly across water sources and mainly depend on the distance from the location of pollutant discharge.For example, PFAS concentrations in water range from 26 to 5200 ng L À1 ; 97,98 for illustration, the lower limit corresponds to approximately one drop in an Olympic swimming pool full of water.Freshwater sources are typically contaminated with PFAS at levels on the order of hundreds of nanograms per liter, whereas concentrations in marine water are at tens of nanograms per liter. 98oncentration techniques.Halogenated organic compounds have been separated from water by nitrifying fluidized-bed biomass 99 or by distillation, taking advantage of volatility differences between organohalogens and water before discharging the water into sewage systems. 100Adsorption and reverse osmosis separation methods have been described to concentrate chlorinated and brominated organic pollutants from dilute natural water sources.
Adsorption techniques.Adsorption utilizes physical or chemical interactions between the surface of a solid (adsorbent) and a solute (adsorbate) to remove pollutants from water sources.In contrast, absorbents are porous materials that take up matter into spaces within and throughout the material.Adsorbable organic halogens are removed from water by adsorption methods, 101,102 such as Pd/Fe bimetallic particles to treat activated sludge of chemical dyestuff wastewater. 103ranular activated carbon and powdered activated carbon are the most frequently used adsorbents, but activated alumina and zeolites have also been employed. 104Two common chlorinated groundwater pollutants, tetrachloroethylene and trichloroethylene, have shown significant adsorption on granular activated carbon. 105Likewise, absorbent materials that are capable of removing brominated water pollutants include metal organic frameworks, 106 biochar, 107 and microplastics. 108verse osmosis.Reverse osmosis is a separation technique that passes water through a semi-permeable membrane that discharges a treated stream (permeate) and a rejection stream (concentrate).Reverse osmosis decreased the concentration of 2,3,7,8-substituted polychlorinated dibenzo-p-dioxins (PCDD), dibenzofurans (PCDFs), and dioxin-like PCBs in water. 109estruction techniques.Destruction of organohalogens is preferable over standalone separation.Separation pre-treatment procedures are often used to increase organohalogen concentrations in water to enhance degradation efficiency [110][111][112] by overcoming substrate mass transport limitations that are inherent with dilute solutions.4][115] Reported destruction methods for pollutant

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This journal is © The Royal Society of Chemistry 2023 remediation include electrochemical, photochemical, mechanochemical, thermochemical, and advanced oxidation processes, or combinations of these approaches. 27The exceptionally high thermodynamic stability of carbon-halogen bonds in organohalogens poses the most significant challenge for destruction efficiency; C-X (X = F, Cl, or Br) bond dissociation energies of common organohalogens are in Table 1.
The bond dissociation energy values, which are a measure for thermodynamic bond strength, show that halogen atoms with higher electronegativity form stronger bonds to carbon atoms (Table 1).Fluorine is the most electronegative element in the periodic table and ergo C-F bonds are the strongest covalent bonds known in organic chemistry.In general, electronegativity of halogens follows F 4 Cl 4 Br, and likewise the trend in bond strengths follows C-F 4 C-Cl 4 C-Br.The C-F bond dissociation energies in PFAS depend on the extent of fluorination of carbon atoms, the position of C-F bonds, and the nature of the head group in the molecule (Fig. 3).
In contrast, in chlorinated and brominated organic compounds, dissociation energies of C-X (X = Cl or Br) bonds depend on the degree of substitution at the carbon of the C-X bond.Thus, the strongest C-X bonds of chlorinated and brominated compounds are found in halogenated phenyls. 124reshold limit values and response levels of common halogenated organic water pollutants Threshold limit values (TLVs) serve as important benchmarks for monitoring workplace exposure.These values denote permissible airborne concentrations of chemical substances that the majority of workers can encounter in a repetitive manner throughout their career span, without detrimental health consequences.The American Conference of Governmental Industrial Hygienists (ACGIH) formulates TLV-timeweighted averages (TLV-TWAs) to safeguard against prolonged exposures, as well as short-term exposure limits to shield against sudden spikes in exposure.It is important to note that TLVs are not designed to establish a rigid demarcation between safe and hazardous exposures.Instead, their primary objective is to safeguard workers from potential health impacts.125 The notification level is the concentration of a hazardous substance in the environment, above which specific actions or notifications are required.Notification levels are often used to assess water pollutants because they represent human health-based advisory levels for chemicals in drinking water, which lack maximum contaminant levels.126 Human health-based advisory levels establish the contaminant concentration in drinking water, below which no adverse health effects and/or aesthetic impacts are expected during specific periods of exposure.127 Actions might include reporting the presence of the substance to regulatory authorities or taking steps to mitigate the exposure or contamination.Notification levels are typically set by regulatory agencies and are designed to alert relevant parties when a certain level of contamination has been reached.In contrast, the response level represents a concentration of a hazardous substance in the environment that triggers specific actions or interventions to mitigate risks and protect human health or the ecosystem.Response levels are set to guide the appropriate measures to be taken when contamination reaches a certain level.These measures might include evacuation, containment, or clean-up, to prevent further exposure or harm.126 TLVs and notification levels of common halogenated organic water pollutants that are discussed in this article are in Table 2.

Destructive remediation of chlorinated and brominated organic water pollutants
Destructive methods are necessary to treat waste streams or sludges after concentration of target pollutants by separation processes.Chlorinated and brominated organic pollutants can be destroyed by incineration, 157 non-thermal plasma discharge, 158 g-irradiation, 159 biological, catalytic, photolytic, photocatalytic, or photochemical processes.Incineration.Incineration has been widely used to reduce the volume, potential infectious properties, and potential toxicity of waste. 160This process burns the hazardous waste to carbon dioxide and water vapor, while also producing byproducts that are released as exhaust gas or remain as solid ash or soot. 160Incineration has the advantage that large amounts of waste can be treated in a short period of time, 160 while additionally offering the possibility of providing heat energy to local communities; 161 however, incineration of halogenated pollutants does not always result in complete mineralization, which can give rise to the release of toxic emissions. 157PCBs can be decomposed thermally; ergo, incineration at high temperatures can be employed for PCB destruction.PCBs are often formed as byproducts of lower-temperature municipal solid waste incineration. 157on-thermal plasma and c-irradiation.Non-thermal plasma treatment utilizes high-voltage electrical pulses to generate a corona discharge that excites electrons in the air to produce singlet oxygen, which reacts with water to generate ozone and hydroxyl radicals for pollutant destruction. 162Gammairradiation utilizes g-rays with sufficient energy, often from a 60 Co source, to ionize atoms, resulting in cleavage of molecular bonds, such as water, to create reactive radical species for pollutant destruction. 162In addition to the inherent safety precautions and high-risk handling procedures of g-rays, the efficiency of g-irradiation relies on the extent which the radiation source has decayed, as that determines the deliverable dose.Non-thermal plasma has the advantage that the efficiency relies on electromagnetic generation of particles, which does not decay over time; however, its high energy demand creates inherent barriers for global use. 163,164Non-thermal plasma discharge has been described for the degradation of PCB-77 (80% within 2 min) using a dielectric barrier discharge nonthermal plasma, 158 whereas PCB-47 was degraded by 70% within 60 min using pulsed corona discharge. 165PCBs can also be decomposed with ionizing radiation in aqueous micellar solutions. 166However, g-radiation hazards originating from the required 60 Co g-irradiation source must be mitigated. 167hemical and biological techniques.Chemical and biological techniques make use of chemicals or microorganisms, respectively, to break down organic matter.Chemical processes are attractive destruction methods due to their simple operation; however, stoichiometric use of chemicals and associated production of sludge create large amounts of chemical waste, contributing to operational cost and complexity of separation. 168,169Biological techniques have the advantage of being self-sustaining processes, which result in less cost compared to methods requiring chemical additives or extensive maintenance.Nevertheless, biological processes require extended time for microbial growth, and the performance can suffer from inherent environmental factors, such as temperature or water composition. 169,170Catalytic or oxidative processes have been reported for breakdown of chlorinated and brominated organic pollutants. 171,172Biological processes have been applied to treat brominated water pollutants, such as anaerobic-aerobic processes for microbial degradation of tetrabromobisphenol-A. 173lectrochemical techniques.Electrochemical techniques utilize electrical energy to induce the mineralization of dissolved contaminants in water.Electrochemical systems offer many advantages, including operation at ambient temperature and pressure, no required auxiliary chemicals, and small footprint, making it an attractive technique for delocalized water treatment. 174However, electrochemical technologies must be improved with regard to toxic by-products that result from inefficient mineralization, and with respect to electrode costs to enable adoption on an industrial scale. 174Chlorinated and brominated organic water pollutants have been degraded by electrochemical processes. 25,175,176Electrooxidative 176,177 and electroreductive 19,178,179 methods were used via direct electrocatalytic electron transfer or indirect interactions with electrochemically produced highly reactive transient species.][182] Photo-assisted techniques.Photo processes and photoassisted techniques utilize light absorption by molecules to directly dissociate target species or indirectly create reactive radical species in solution that carry out the degradation.Photo processes have several advantages, including operation at ambient temperature and pressure, low energy requirement, and no need for additional chemicals. 27Nevertheless, current photo processes lack degradation and mineralization efficiency, resulting in by-product formation. 27Methods involving illumination with light in the visible to vacuum ultraviolet range, such as direct photolysis or indirect photocatalytic and photochemical processes have been utilized for chlorinated and brominated pollutant mineralization. 183,184onolysis.Sonolysis utilizes the compression and expansion cycles of ultrasound waves that produce hot spots with high temperatures (approximately 5000 K) and pressures (approximately 1000 atm), due to cavitation bubble collapse. 185In these hot spots, molecular bonds are cleaved directly through pyrolysis or indirectly through reactions with reactive radical species produced via pyrolysis. 185Sonolysis is advantageous because no chemical additives are required and it is a simple process; 186,187 however, the localized production of high concentrations of reactive radical species is limited by recombination leading to inefficient degradation. 1880][191] To enhance the degradation efficiency of sonolysis processes, auxiliary chemicals have been added to solutions that contained chlorinated or brominated pollutants. 189,192cent progress in the destruction of chlorinated organic water pollutants PCBs.Multiple PCBs were degraded via a combination of adsorption, photodegradation, and heterogeneous Fenton oxidation reactions, using a multifunctional magnetic bcyclodextrin/graphitic carbon nitride catalyst (Fe 3 O 4 @b-CD/g-C 3 N 4 ), with a degradation efficiency for different PCBs in wastewater ranging from 77% to 98%. 193PCBs from transformer oil were photocatalytically degraded using carboxymethylb-cyclodextrin modified Fe 3 O 4 @TiO 2 , with a degradation rate of PCBs of 83% after 16 minutes. 194Dielectric barrier discharge non-thermal plasma degraded PCB77 in aqueous solution with a removal efficiency of PCB77 of 80% with helium as discharging gas and approximately 75% with oxygen as discharging gas. 158ioxins.Dielectric barrier discharge in a lab-scale reactor degraded 2,3,7,8-TCDD in fly ash with a removal efficiency of 92%; this process was also used for degradation of other PCDD/ Fs-containing fly ash, whose degradation efficiency depended on input energy and discharge time. 195Another reported method to degrade 2,3,7,8-TCDD is use of extracellular fungal ligninolytic enzymes that were made of laccase enzymes.Ligninolytic fungus Rigidoporus sp.FMD21 degraded 2,3,7,8-TCDD by 77.4% in 36 days and produced 3,4-dichlorophenol. 196etrachloroethene (perchloroethylene, PCE).Degradation of PCE occurred under aerobic conditions using Sphingopyxis ummariensis bacteria in a gas-recycling fixed-bed bioreactor. 197This process was more efficient with lower concentrations of PCE and achieved complete degradation of PCE in 25 hours. 197PCE degradation using cobalt-mediated electroscrubbing with boron-doped diamond (BDD) coating supported on silicon or tantalum substrate anodes and stainless steel cathodes was reported. 198The process worked by volatilizing of liquid PCE, followed by the treatment.BDD on silicon substrate anode reached a PCE removal efficiency of 75.7% in 2 hours, while BDD on tantalum substrate achieved a PCE removal efficiency of 90.5% in 2 hours. 198Complete degradation of PCE was reported at pH 7 in 4 hours using nano-magnetite catalyzed with glutathione, with oxalic acid as the major byproduct. 199richloroethene (TCE).Degradation of TCE by sodium percarbonate activated with Fe(II)-citric acid complex in the presence of surfactant Tween-80 has been described.At optimal conditions, 93.2% degradation efficiency was achieved in 15 minutes. 200TCE was degraded using nanoscale calcium peroxide activated by Fe(II)/FeS, which enhanced generation of hydroxyl radicals, achieving 99.5% TCE removal efficiency in groundwater. 201Polyvinyl alcohol coated nano calcium peroxide activated by Fe(II)/FeS or Fe(III)/FeS has been used to degrade TCE, with maximum degradation of 91% and 95% for Fe(II)/FeS or Fe(III)/FeS activated polyvinyl alcohol coated nano calcium peroxide, respectively. 202Sequential anaerobic and aerobic treatment in the presence of the cyclic ether stabilizer 1,4-dioxane degraded TCE.The anaerobic treatment used halorespiring consortium SDC-9 and effectively removed TCE, forming vinyl chloride (VC) and cis-dichloroethene (cis-DCE) as byproducts.These by-products were removed along with the 1,4dioxane during the subsequent aerobic bioaugmentation with Azoarcus sp.DD4. 203ther chlorinated organic water pollutants.Bimetallic zerovalent iron nanoparticles have been reported to degrade several chlorinated organic compounds, such as vinyl chloride (VC), 1,2-dichloroethene (DCE), TCE, and PCE.Bimetallic zero-valent iron nanoparticles with palladium and with nickel have completely degraded all compounds within 24 hours.VC, DCE, and TCE were completely degraded in 2 hours.PCE was degraded by about 97% and about 89% in 4 hours using Pd or Ni modified zero-valent iron nanoparticles, respectively, and complete degradation was achieved in 24 hours for both materials. 204egradation of chlorinated volatile organic compounds from contaminated groundwater was achieved by an O 3 -bubble column reactor with a carrier-bound TiO 2 /ultraviolet light system, with degradation efficiencies of 98% for cis-1,2-DCE, TCE and PCE and of 85% for trichloromethane without detectable by-product formation. 205TCE and cis-1,2-DCE were degraded with Cupriavidus sp.CY-1 bacteria, whose growth was supplemented with TCE or cis-1,2-DCE and phenol or Tween 80 as a co-substrates.Use of CY-1 bacteria, whose growth was augmented by phenol, TCE and cis-1,2-DCE were converted into poly-b-hydroxybutyrate (PHB), which is a biodegradable plastic. 206cent progress in the destruction of brominated organic water pollutants Tetrabromobisphenol-A (TBBPA).Peroxymonosulfate in aqueous solution was activated by Ce, Sn, or Sb doped copper ferrite, CuFe 2 O 4 , catalysts prepared by a sol-gel combustion method, to degrade TBBPA. 207A TBBPA removal efficiency of 90.1% in weakly basic conditions was achieved with Sb-doped CuFe 2 O 4 . 207In another study, dielectric barrier discharge was used to completely decompose TBBPA in wastewater in 12 minutes, forming phenol, bisphenol A, catechol, hydroquinone, and 3,5-dibromophenol as by-products. 208Bimetallic Co/Fe metal-organic frameworks/cellulose nanofiber membrane as a catalyst in a sulfate radical advanced oxidation process activated peroxymonosulfate and completely degraded TBBPA in 30 minutes at optimal conditions. 209,4,6-Tribromophenol (TBP).An in situ peroxymonosulfate oxidation process with added chloride completely degraded TBP in salty wastewater, albeit with formation of undesired persistent halogenated products. 210Ultraviolet photolysis of TBP in the presence of hydroxylamine achieved a debromination rate of 89.9% in 1 hour. 211exabromocyclododecane (HBCD).Near-complete aerobic biodegradation of aqueous HBCD was obtained by Rhodopseudomonas palustris YSC3 strain at 35 1C and neutral pH, forming bromide ions, pentabromodyclododecanol, and pentabromocyclododecene as by-products. 212Isotope-labeled [ 13 C]-HBCD was efficiently mineralized in 5 days into 13 CO 2 using organic montmorillonite-supported nanoscale zero-valent iron coupled with the bacterial strain Citrobacter sp.Y3. 213 HBCD removal and mineralization was obtained by an ultrasound-based advanced oxidation process, which completely degraded HBCD and accomplished 72% of total organic carbon removal in 40 minutes. 191Complete degradation of HBCD was observed using nanoscale zero-valent aluminum in 8 hours in an ethanol/water solution at 25 1C, producing completely debrominated cyclododecatriene with 67% yield. 214A ball-milled aluminum-carbon composite has been prepared to enhance the absorption and degradation of HBCD, completely absorbing HBCD in water in 1 hour and debrominating 63.44% of the pre-sorbed HBCD in 62 hours. 215We note that zero-valent firstrow transition metals or aluminum may oxidize in ambient aqueous conditions that are needed for global scalability.
2,2 0 ,4,4 0 -Tetrabromodiphenyl ether (BDE-47).Reduction at zero valent zinc with cetyltrimethylammonium chloride surfactant achieved a BDE-47 removal efficiency of 98.6% in 1 hour, followed by a Fenton oxidation that decomposed all obtained debromination products into short-chain carboxylic acids that were mineralized in 2 hours. 216Complete degradation of BDE-47 was achieved in 3 hours with a thermally activated persulfate system, forming one low-toxicity oxidation product. 217A photocatalytic process using Ag/TiO 2 was developed for the degradation of BDE-47 in Triton X-100 surfactant solution under anaerobic conditions, predominantly producing diphenyl ether and the harmful bromodiphenyl ethers BDE-28, BDE-15, BDE-3, 218 which were found to be phytotoxic, 219 or exhibited hepatic 220 or reproductive toxicity 221 in mice.BDE-47 was degraded using a Fe(II)-catalyzed peroxymonosulfate activation process with the addition of gallic acid to accelerate the cycling of Fe, which enhanced peroxymonosulfate activation, reaching a degradation efficiency of 85% in 72 hours. 222A functional bacterial consortium QY2 with an addition of methanol to enhance degradation efficiency and accelerate the debromination, hydroxylation, and phenyl ether bond breakage of BDE-47 completely removed BDE-47 in 7 days. 223

Destructive remediation of fluorinated organic water pollutants including PFAS
The destructive remediation of fluorinated organic compounds, particularly PFAS, is an area of intense research.Methods for defluorination of the common PFAS chemical perfluorooctanoic acid (PFOA) have been reported by electrochemical reduction at a Rh/Ni cathode in dimethyl formamide via hydrodefluorination, 224 hazardous g-irradiation with a 60 Co source in an alkaline solution under N 2 -saturated conditions, 159 and mineralization of perfluorocarboxylic acids, including PFOA, via the formation of rapidly decomposing carbanions in polar, aprotic dimethyl sulfoxide electrolyte; the carbanion-based mechanisms can only operate in waterfree, polar, aprotic solvent and fail to degrade sulfonic acid PFAS, such as PFOS. 225Further, 19 F-NMR was used to quantify PFOA defluorination, which appeared to have a detection threshold of Z5 mM (approx.2000 ppm PFOA). 225

Globally scalable, viable technologies must work in aqueous media and
This journal is © The Royal Society of Chemistry 2023 enable destruction of PFAS with much lower concentrations.Perfluorocarboxylic acids, such as perfluoro-butanoic, pentanoic, hexanoic, heptanoic, and octanoic acid (PFOA), were mineralized using a Ce-doped nanocrystalline PbO 2 film electrode; 226 the toxicity of lead poses challenges.While these reports are mechanistically intriguing, only cost-and energysaving aqueous methods that utilize nontoxic, nonprecious materials and renewable electricity will be viable and sustainable on a global scale.Electrooxidation of PFOA and PFOS at Magne ´li-phase Ti 4 O 7 ceramic anodes outperformed mineralization at Ce-doped PbO 2 and Ti-modified boron-doped diamond electrodes, due to faster oxidation rate. 227PFOA in water was degraded by ultraviolet-visible light assisted Zn x Cu 1Àx Fe 2 O 4oxalic acid system, using a ferrite-based catalyst that allowed for magnetic catalyst recovery. 228echnoeconomic analyses of destruction techniques are indispensable to assess viability on a global scale.Significant research has been done to determine the most efficient destruction techniques. 27Energy efficiency is important to lower operational costs and improve carbon footprints, whereas capital expenditures matter for assessing the affordability of units.Estimated energy and capital cost requirements of existing PFAS destruction techniques, based on literature data, 167,[229][230][231][232][233][234][235][236][237] are shown in Fig. 4. Capital expenditure values refer to cost of equipment in the United States, are given in U.S. dollars, and are needed to treat at least one cubic meter of polluted water.
Incineration requires an initial investment of $41 938 050 for a one-line ''turn-key'' incineration plant. 83The average amount of energy required for incineration is 0.45 kW h m À3 in the U.S., which has been measured and varies depending on the country and the type of incineration plant used. 233The reported initial investment cost of g-irradiation is $4 176 150, which includes the sum of the cost of the electron beam accelerator with lifespan of 15 years and personnel costs needed to run a girradiation facility. 238The equivalent electrical energy requirement to achieve 90% sulfadiazine degradation by g-irradiation from a 60 Co source, taken here as a proxy for PFAS degradation, at a constant dose rate of 6.69 kGy h À1 has been reported to be 18 kW h m À3 . 239The capital investment for non-thermal plasma 231,237 of $92 329 was obtained from a capital cost approximation for industrial wastewater plants. 163,240The energy required to achieve a high removal rate of PFOA from water using non-thermal plasma is 100 kW h m À3 , which was deduced from the energy efficiency of non-thermal plasma setups. 237Sonolysis by ultrasonication has a capital investment cost of $9 390 000, which has been calculated to include the part replacement cost, labor cost, analytical costs, chemical costs, and electrical costs. 230The energy requirement for ultrasonication (1475 kW h m À3 ) was calculated using the energy efficiency of ultrasonication in g (kW h) À1 . 235,241hemical oxidation of PFOA by stoichiometric amounts of permanganate, 242 hydrogen peroxide, or persulfate 26,243 is costprohibitive and creates large amounts of chemical waste.Methods based on electrochemical processes have gained popularity because of lower energy demands than physical destruction methods, i.e. incineration, g-irradiation, non-thermal plasma, and sonolysis.PFAS destruction by supercritical water oxidation 244 requires high initial investment; 245 we were unable to find numbers for investment costs.The thermal energy requirement for an efficient supercritical water generation facility is 5 MW for 250 metric tons of water per day. 2468][249][250][251][252][253][254][255][256][257] The inter-electrode distance matters in electrolyzers because smaller distances are concomitant with less ohmic losses. 258A generally accepted upper limit for inter-electrode distance in aqueous systems is 10 cm, which necessitates 10 m 2 geometric electrode area to treat 1 m 3 of polluted water, 258 rendering boron-doped diamond (BDD) electrodes cost-prohibitive. 259At smaller interelectrode distances, which generally result in higher electrocatalytic performance, the geometric electrode area requirement increases for treatment of a 1 m 3 batch, making the economics of BDD electrodes even more unfavorable.The energy required to halve an initial PFOA concentration of 15 mg L À1 with a BDD electrode of 38 cm 2 geometric area and an inter-electrode distance of 4 mm was reported as 180 W h L À1 at 50 mA cm À2 , for the treatment of 250 mL solution. 234Conversion of these numbers to the treatment of 1 m 3 of polluted water requires BDD electrodes of 15.2 m 2 at a cost of $8.58 million 229 and 180 kW h m À3 electrical energy (Fig. 4).Electrochemical degradation of PFOA on ultrananocrystalline BDD coated on niobium electrodes additionally produced toxic perchlorate. 234Ultraviolet-light-assisted electrochemical PFAS defluorination required cost-prohibitive platinum electrodes and N 2 -saturated electrolyte. 260Ergo, potentially viable PFAS destruction technologies must be more cost-effective and energy-saving to achieve economic feasibility and reduce carbon emissions.

Recent progress in the destruction of PFAS
Perfluorooctanoic acid (PFOA) has been degraded anaerobically to shorter chain perfluoroalkyl carboxylic acids and produced Fig. 4 Estimated electrical energy and capital expenditure requirements fluoride in biosolids using Acidimicrobium sp.strain A6 and ferrihydrite. 261PFOA has also been degraded efficiently by 80% over ball-milled boron nitride in deionized water or by 60% in simulated drinking water using photocatalysis for 2 hours. 262he photocatalytic degradation of PFOA under UV-A and UV-C illumination with boron nitride was improved by using a composite of boron nitride and titanium oxide instead of neat boron nitride or titanium oxide.However, complete mineralization was not achieved and only shorter chain perfluorinated carboxylic acids were produced.The composite was capable of degrading PFOA by 68% in 7 hours using natural sunlight. 263FOA was degraded via a photocatalytic process that used a carbon-modified bismuth phosphate composite, absorbing PFOA in 2 hours and achieving nearly complete decomposition in situ in 4 hours under UV irradiation.264 Another composite that has been developed for the photocatalytic degradation of PFOA in water is iron (hydr)oxides/carbon sphere (FeO/CS) composite.This material almost completely adsorbed PFOA in 1 hour, which subsequently underwent photodegradation and defluorination.During this process, PFOA was photodegraded by 95.2% and defluorinated by 57.2% in 4 hours.265 Another photocatalytic process that used UV light and Bi 3 O(OH)(PO 4 ) 2 in acidic conditions has been reported to degrade PFOA.However, this process was unable to degrade shorter chain PFAS unless some changes were made to the reactor system.With this system, challenges for implementation in real waters exist because the reaction was quenched by chloride and sulfate.266 Complete defluorination of PFOA was achieved after 6 hours in water using a dual-frequency ultrasonic activated persulfate.This method can be used for other PFAS compounds, but not as effectively. 267An electrochemical degradation process for PFOA using sodium sulfate has been developed, which enabled 99.5% degradation and 50% fluoride generation after 4 hours.268 A sorptive photocatalyst Fe/ TNTs@AC, which was based on activated carbon and titanium oxide, was synthesized and used for PFOA degradation.This catalyst completely absorbed PFOA in 1 hour, degraded it by 90% in 4 hours using UV radiation, and then mineralized 62% of the degraded PFOA to fluoride.269 Another photocatalytic process that has been developed to degrade PFOA in aqueous solutions used In 2 O 3 nanoparticles.This process worked most efficiently in acidic (pH = 2) conditions and successfully decomposed most of PFOA to fluoride and carbon dioxide within 90 minutes under UV irradiation, while also achieving 95.99% of PFOA defluorination.270 Chemical mineralization of perfluorocarboxylic acids is mechanistically intriguing, 225 but works only in polar, non-protic electrolytes, such as dimethyl sulfoxide (DMSO), hampering global scalability, and fails to degrade sulfonic acid PFAS, such as PFOS.
Perfluorooctane sulfonic acid (PFOS) degradation has been reported, using sonolysis by 96.9%, 93.8%, and 91.2% at 400 kHz, 500 kHz, and 1000 kHz, respectively, in 4 hours. 271nother method to degrade PFOS in an aqueous solution used reverse vortex flow gliding arc plasma; PFOS was degraded by 93.1% in 1 hour.This method can also be used to degrade other PFAS compounds, such as PFOA or PFDA and others. 272Both PFOA and PFOS were defluorinated by 88% in 1 hour and 92% in 24 hours, respectively, using a UV light in a system that contained sulfite and iodide.Adding iodide to the UV light illuminated sulfite system greatly accelerated the defluorination of many PFAS.This system has achieved a complete removal of both PFOA and PFOS from concentrated mixtures in NaCl brine. 273Electrooxidation using titanium suboxide anodes were used to degrade multiple PFAS compounds, such as PFOA and PFOS, whose concentrations decreased very quickly and approached zero in 1 hour. 274Another method to degrade PFOA and PFOS as well as other PFAS compounds is girradiation, using a 60 Co source.PFOS degradation was more efficient with branched PFOS isomers compared to linear molecules.Branched PFOS isomers were degraded almost completely, while PFOA was degraded by 87%. 275A nonthermal plasma generator was custom-built to remove PFOA, perfluorohexanoic acid (PFHxA), and PFOS from water in both ultrapure and groundwater.In 30 minutes, PFOS was degraded completely in ultrapure water and by 85% in groundwater.PFHxA was degraded by 35% in ultrapure water and by 40% in groundwater, while PFOA degradation reached about 50% in both water systems. 276PFOS and PFOA were significantly degraded in soil and groundwater by high dose electron beam technology.This process enabled the decrease of PFOS and PFOA concentrations in groundwater by 87.9% and 53.7%, respectively. 277A duo-functional tri-metallic-oxide hybrid photocatalyst has been developed for the degradation of many PFAS compounds. 278It possessed a high adsorption capacity and achieved 99.8% and 99.4% adsorption efficiency for PFOS and PFOA respectively, and it exhibited a high defluorination rate up to 67.6% for PFOS and 74.8% for PFOA.With this catalyst, PFOS was degraded by 95.5% in 5 hours, while PFOA was degraded by 98.9% in 30 minutes. 278Boron-doped graphene sponge anodes have been reported for the degradation of PFAS compounds by electrochemical oxidation.This method had a higher removal efficiency for longer-chain than for shortchain PFAS, but its defluorination efficiency was lower than that of other methods. 279GenX, which is a short-chain PFAS, was mineralized by an electro-Fenton process that paired a graphene-coated nickel foam electrode with a boron doped diamond electrode.This process achieved 92.2% mineralization after 6 hours of treatment. 280

Aqueous advanced redox processes
Advanced oxidation processes, advanced reduction processes, and combined hybrid processes have emerged as promising strategies in water remediation due to potentially high organohalogen destruction efficiencies. 113,2813][284][285] The advantages of advanced redox processes have led to intense research, such as photolytic, photochemical, photocatalytic, cavitation, electrochemical, and ionizing radiation approaches for water remediation, as well as combinations of these methods. 113

À
), sulfite radical anion (SO 3 À ), 303 and hydrogen radical (H ), as well as non-radical hydrated electrons (e À aq , aqueous lifetime 0.43 or 8.6 ms at pH 7.0 or 9.5, respectively). 113,287,304xygen-free conditions are often impractical, especially since the water oxidation half reaction can produce O 2 , 305,306 membranes that separate oxidation and reduction half reactions cannot completely exclude oxygen crossover, 307 and airtight seals are inherently challenging on a large scale.

Mechanistic aspects of electrocatalytic aqueous advanced redox processes
A quantitative mechanistic understanding of aqueous advanced redox processes is lacking to date, albeit urgently needed to accelerate the development of viable aqueous electrocatalytic organohalogen destruction techniques.Radicals and reactive oxygen species play a major role in aqueous advanced redox processes, and the mechanisms of their formation and reactions with organohalogens must be understood in detail.Electrocatalysts significantly lower energy requirements for electrochemical transformations, 305 especially in aqueous systems, where potential-leveling by proton-coupled electron transfer (PCET) decreases needed energy inputs. 308active oxygen species of the water-oxygen system Aerobic aqueous processes that produce H 2 O 2 , OH, HOO , or OO À are based on the electrochemical transformations of the water-oxygen system (Fig. 5).Most of the transformations are electrochemical oxidations or reductions that proceed through energy-saving, potential-leveling PCET steps, which require that equal numbers of protons and electrons are transferred. 308eactive oxygen species that are relevant for organohalogen destruction, i.e.H 2 O 2 , OH, HOO , and OO À , can be produced directly or via H 2 O 2 decomposition.Water oxidation can only produce O 2 , H 2 O 2 , or OH through direct electrocatalysis.The one-electron-one-proton or two-electron-two-proton water oxidation reactions, given here together with standard potentials E 0 in V vs. the standard reference electrode (SHE), i.e.H 2 O -OH (aq) + (H + + e À ), E 0 = 2.38 V SHE , or 2H 2 O -H 2 O 2 + 2(H + + e À ), E 0 = 1.76 V SHE , are kinetically easier than the four-electron-four-proton reaction, i.e. 2H 2 O -O 2 + 4(H + + e À ), E 0 = 1.23 V SHE , but thermodynamically more uphill. 3Molecular oxygen can be reduced to H 2 O 2 , OH, HOO , or OO À (Fig. 5). 309Electrochemical reductions and oxidations that involve the transfer of less electrons and protons are kinetically faster than those that require the movement of more electrons and protons in the H 2 O-O 2 -system (Fig. 5).The electronic and chemical structures of intermediates of one-electron-one-proton transformations are known from computational density functional theory work, which cannot accurately capture concerted two-electron steps. 2,3,6Together with the thermodynamic potentials below, these kinetic considerations allow as a function of electrolyte pH estimates which reactant species are likely formed in the highest concentrations for efficient destruction of halogenated organic water pollutants.
Hydrogen peroxide is a precursor for OH, HOO , or OO À radical generation (Fig. 5) and, therefore, often called a pre-or pro-radical species.1][312][313] Addition of H 2 O 2 to the electrolyte was found to quench the production of toxic perchlorate at BDD electrodes. 314

Thermodynamic potentials of reactive oxygen species
The thermodynamic potentials of reactive oxygen species determine their oxidation or reduction strengths in AOPs and ARPs of halogenated organic water pollutants.The protonation state and thermodynamic potential of aqueous reactive oxygen species depend critically on the electrolyte pH (Table 3).For example, at pH 0, OH and HOO are strong aqueous oxidants with standard potentials (i.e.thermodynamic potentials at pH 0) of 2.73 and 1.66, respectively, and both radicals possess sufficient thermodynamic driving force for PFAS oxidation.In contrast, at pH 14, the deprotonated hydroxyl radical anion (O À ) possesses a thermodynamic potential of 1.77 V, whereas the thermodynamic potential of the superoxide radical anion (OO À ) is only 0.65 V, which is insufficient to oxidize PFAS.The protonation state of chemical species depends on their pK a value and the pH of the aqueous solution; if the pH value is larger than the pK a value of a species, the molecule is deprotonated.Aqueous pK a values of reactants and chemicals that are relevant for PFAS destruction are shown in Table 4.

Mechanistic role of oxygen radicals for PFAS destruction
Reactive oxygen radicals, such as OH, O À , HO 2 , OO À , are key species in AOPs and ARPs for C-C and C-F bond cleavage. 113,322Superoxide radical anions (OO À ) can oxidize PFAS molecules; OO À can also act as a reductant and is the predominant species at pH 4 5, as HOO has an aqueous pK a of 4.8. 309Hydroxyl radicals in aqueous solution are widely considered insufficient for PFOA or PFOS degradation 310,323,324 because OH radicals have an approximately 10Â shorter aqueous diffusion distance than OO À radicals. 325,326Quenching experiments with saturated alcohols that selectively scavenged solution OH radicals had no effect on the degradation of PFAS. 327,328The exact role of O(H) radicals in aqueous electrocatalytic PFAS destruction is still debated.Recently, decreased degradation efficiency was observed in the presence of unsaturated allyl alcohol that is capable of quenching OH radicals in the vicinity of the anode, suggesting that adsorbed OH radicals played a significant role in PFAS degradation. 328,329Adsorbed OH radicals are physisorbed; analog surface-bound species are non-radical hydroxide anions (OH À ) that lack the thermodynamic driving force for PFAS destruction. 309xiliary radicals and non-radical species Addition of suitable solutes to aqueous media enables the formation of auxiliary radicals and non-radical redox agents in aqueous AOPs and ARPs.1][332] Addition of iodide, dithionite, or ferrocyanide to anaerobic electrolytes can aid the detachment mechanism to produce hydrated electrons. 260,333Dissociation of dithionite will produce reductive sulfur dioxide radicals. 302ulfite assists PFAS degradation through formation of persulfite. 323,334Stoichiometric PFOA oxidation by permanganate has been observed. 242Borate is known to react with H 2 O 2 , to produce peroxoborates that are stable at pH 8 to 12 and highly reactive towards nucleophiles. 3357][338] In the case of perfluoro-compounds, such as PFOA and PFOS, that do not contain carbon-bound hydrogens and CQC double bonds, a different initial oxidation step is required that starts with the elimination of the head group in photo-or electrochemical processes. 282,339,340An electron transfer from the carboxylic acid to the anode creates a carboxyl radical that undergoes decarboxylation to produce the carbon-centered perfluoroalkyl radical ( C n F 2n+1 ) and CO 2 . 176,183,274Likewise, perfluorinated sulfonic acids are desulfonated via an analogue initial electron transfer. 256,282,341,342The generated carbon-centered radicals quickly react with surrounding dissolved oxygen, water, or other radicals to form smaller carbonyl species.Once initiated, subsequent propagation reactions between the parent and daughter species of the decomposing pollutant with highly reactive, strong oxidants will ultimately lead to complete mineralization of perfluorinated compounds. 286eductive dehalogenation via hydrated electrons that are generated by ultraviolet photolysis from sulfite requires oxygenfree conditions, which are impractical on a large scale.Added sulfite can effectively deoxygenate aqueous solutions, forming sulfate, which in turn can form SO 4 À for AOPs.Sulfite promotes the formation of SO 3 À , which is a mild oxidant 343 that reacts rapidly with oxygen to produce SO 5 À and subsequently SO 4 À and OH radicals, which are stronger oxidants and can more effectively degrade organohalogen pollutants. 301

Electrocatalytic aqueous generation mechanisms of reactive species
Electrocatalytic AOP and ARP mechanisms proceed through direct electrocatalysis at materials surfaces, 96,115,283,344 indirect solution reactions, 176,345 or assisted reactions at materials and in solution. 332,346,347In aqueous electrolyte without auxiliary solutes, aerobic electrocatalysis encompasses anodic water oxidation to H 2 O 2 (precursor for AOPs), OH radicals (for AOPs), or O 2 (loss process), 3 and cathodic oxygen reduction at suitable catalysts 348,349 to H 2 O 2 , OH or OO À radicals, or water; H 2 O 2 reduction to water can also occur (see Fig. 5 and  6A).Produced reactive species can diffuse into the bulk solution if they are sufficiently long-lived (see aqueous advanced redox processes) and react in indirect reactions to produce additional reactive species.The aqueous lifetime of OH is 0.02 ms, whereas that of OO À is 1.3 ms, resulting in a diffusion distance of OO À radicals (30 nm) that is approximately 10Â longer than that of OH radicals (4.5 nm). 325,326The diffusion distances of both OH and OO À radicals are too short for diffusion from electrodes into the bulk solution, but solution OH and OO À radicals can be created by decomposition of H 2 O 2 in strongly acidic or alkaline water. 350,351In the absence of dioxygen, i.e. in anaerobic media, anodic water oxidation produces OH radicals, H 2 O 2 , or O 2 , which requires deoxygenation of the aqueous electrolyte, e.g. by sulfite, 343 to ensure completely oxygen-free conditions.Direct, unassisted cathodic electrocatalysis in anaerobic electrolyte consists of H 2 O 2 reduction to water (Fig. 6B).Deep-ultraviolet light, 352 radiolysis, 297 or sonolysis 297 352 or OH or OO À radicals by radiolysis 297 or sonolysis 297 processes that assist electrocatalysis (Fig. 6C).Assisted anaerobic electrocatalysis produces OH radicals, H 2 O 2 , or O 2 by anodic water oxidation (Fig. 6D).Elimination of O 2 to ensure anaerobic conditions and immediate decomposition of H 2 O 2 to OH radicals by deep-ultraviolet light, 352 or OH or OO À radicals by radiolysis 297 or sonolysis 297 leave only OH radicals (or OO À radicals if radiolysis or sonolysis were used) as reducible species for the cathode half reaction (Fig. 6D); the highly energetic assisting processes continuously generate OH (or OO À ) radicals in the entire electrochemical cell, including in the vicinity of cathode.Assisted anaerobic electrocatalysis (Fig. 6D) additionally enables the production of hydrated electrons, e À aq , and hydrogen radicals, H , which are needed for ARPs, and which are only accessible in completely O 2 -free aqueous electrolyte.
The aqueous generation mechanisms of reactive oxygen species and radicals are highly entangled, creating complex reaction networks.Much research has been dedicated to the elucidation of mechanisms.Pollutant degradation via AOPs and ARPs involves three principal steps: (i) reactive species generation, (ii) initial attack on the pollutant, and (iii) subsequential attacks on the pollutant until mineralization is complete. 113Direct reactions occur at the electrode surface through electron transfers between the electrode and the chemical substrate. 283In electrochemical oxidation reactions, the anodically generated holes must have sufficient electrochemical potential to create oxidizing agents, without turning on the four-electron-four-proton water oxidation electrocatalysis to dioxygen.Suppression of O 2 evolution requires applied potentials at or below the oxygen evolution potential, which is comprised of the thermodynamic potential and kinetic overpotential at the chosen pH conditions and catalyst materials. 96,305,306Direct oxidation reactions are often slow because substrate adsorption at the anode controls the reaction rate.Due to these slow kinetics and limited useful applied anodic potential, direct electrocatalytic reactions typically do not result in complete mineralization of pollutants. 96uasi-direct redox reactions occur at the electrode-electrolyte interface via physisorbed or chemisorbed redox species, 115,344 which are typically generated through reactions between the electrode and the aqueous electrolyte.Thus, the oxidizing strength in these processes is governed by the thermodynamic potential of the produced reactive species.Quasidirect reactions have inherent mass transport limitations to and from the bulk solution because these processes must take place in the vicinity of the electrode-electrolyte interface.
Indirect processes occur in the bulk electrolyte, to where mediators migrate after electrochemical generation at the electrode-electrolyte interface, to react with pollutant species. 176,345Reactive species for indirect AOPs and ARPs are typically long-lived to enable long diffusion distances into the bulk solution.Therefore, reactive oxygen species generated at the electrode-aqueous electrolyte interface are of limited use in these reactions due to their proclivity for recombination and short lifetimes despite being among the strongest oxidizing agents. 353Some halogenated redox agents have longer lifetimes in comparison to those of reactive oxygen species, however, their selective reactivity and lower oxidation strength limits their reaction efficiency in the bulk electrolyte.Other advanced redox species, like sulfate radicals, possess longer lifetimes and simultaneously similar or greater oxidation strengths compared to reactive oxygen species, enabling indirect redox reations. 165,354,355ndirect processes have been coupled with other advanced redox activation processes to further increase the degradation efficiency of electrochemical systems. 332,346,347,356Ultraviolet or visible light irradiation of electrochemical systems has most often been used.These coupled processes maintain similar reaction networks as dark electrochemical systems, with photoactivation of electrochemically generated mediators, so that increased concentrations of redox agents are produced at the electrode surface and within the bulk electrolyte. 332,346,347hotochemistry is frequently used to destroy pollutants and can be classified as three processes: photolysis, photochemical, and photocatalysis.Photo-assisted AOPs and ARPs have similar advantages as electrochemical processes, such as operation at ambient temperature and pressure, low operating costs, and no generation of waste streams.
Photolysis is the direct absorption of light by chemical substrates for direct degradation via homolytic bond scission or direct light absorption by water to produce highly reactive redox agents for indirect degradation. 113The mechanism of photolysis consists of three steps: (i) light absorption that excites electrons in the molecule, (ii) primary photochemical processes that transform photoexcited molecules or result in relaxation back to the ground state, and (iii) secondary thermal reactions that transform the intermediates that were produced in step (ii). 357Photolysis is limited to pollutants that exhibit large molar absorption cross sections and quantum yields, which restricts the overall applicability of photolysis. 358Previous research has demonstrated applications of photolysis for pollutant degradation, [359][360][361] but photolysis is mainly used for the inactivation of pathogenic microorganisms. 362,363irect ultraviolet photolysis of anaerobic water to form OH radicals, H atoms, and hydrated electrons, e À aq , (see Tables 3  and 5 for thermodynamic potentials) has been demonstrated with vacuum ultraviolet (VUV) wavelengths o200 nm, albeit with low quantum yields and very slow breakdown of PFAS. 364,365Use of VUV irradiation is impractical because of the high absorption cross section of most materials at wavelengths below 200 nm.Deep ultraviolet irradiation at 254 nm is inefficient for direct photolysis of PFOA. 366 ) via radical generation and indirect pollutant oxidation. 323,324,367ddition of transient chemical oxidants to water can overcome photolysis challenges and enhance the overall degradation efficiency of persistent pollutants 95,368 in a process known as photochemical degradation.Mechanistically, photochemical This journal is © The Royal Society of Chemistry 2023 processes proceed through similar pathways as photolysis, i.e., photoactivation of transient chemicals to create highly reactive species, which then interact with pollutants and other molecules in the surrounding bulk electrolyte solution, creating complex reaction networks. 369Typical oxidants in aqueous photochemical organohalogen destruction processes are H 2 O 2 , ozone, peroxosulfate, peroxomonosulfate, and sulfite. 113hotocatalysis employs a semiconductor catalyst to lower the activation energy required to photoactivate water and initiate AOPs and ARPs for pollutant degradation.Semiconducting photocatalysts for oxidation reactions must be n-type to take advantage of the band bending at the semiconductorelectrolyte interface to extract photogenerated holes that can perform oxidation reactions; conversely, photocatalysts for reduction reactions must be p-type to enable enhanced injection of photogenerated electrons into reducible species at the semiconductor-electrolyte interface.Band bending occurs under equilibrium conditions at the junction between a conductor and a semiconductor, [370][371][372][373] here the conducting electrolyte and the solid-state semiconductor photocatalyst.When a conductor and semiconductor are in contact, free electrons will transfer between the conductor and semiconductor because of the work function difference, to align the Fermi levels of both materials.The Fermi level, i.e., the total electrochemical potential of electrons, can be considered as the hypothetical energy level of an electron.Under equilibrium conditions, a Helmholtz double layer forms at the conductorsemiconductor interface, where the conductor and semiconductor carry opposite charges near their surfaces due to electrostatic induction.A charge imbalance arises because semiconductors have a low concentration of free charge carriers; therefore, the electric field at conductor-semiconductor junctions cannot effectively be screened in the semiconductor, which causes the free charge carrier concentration near the semiconductor surface to be depleted relative to the bulk.This interfacial region is called the space charge region. 374In n-type semiconductors, the Fermi level is closer to the conduction band than the valence band, and the electron concentration is larger than the hole concentration.Electrons are the majority charge carriers.Therefore, electrons are depleted in the space charge region, leading to excess positive charges, i.e. photoholes that can perform oxidations at the interface.In the space charge region, the energy band edges in the semiconductor are continuously bent upwards if the semiconductor work function is smaller than that of the adjoining conducting medium, i.e. the electrolyte.This happens due to the charge transfer induced by the electric field at the junction, and the effect is called band bending.Besides different Fermi levels at a semiconductor junction, an external electric field, adsorbed species, or surface states (due to termination of lattice periodicity of a material at the surface) can also induce band bending near the semiconductor interface. 374Band bending can significantly decrease detrimental electron-hole pair recombination rates and enhance carrier transport to the semiconductor surface, 374 where redox reactions occur.
The general mechanism of photocatalysis starts with adsorption of light with a photon energy equal to or greater than the bandgap of the semiconductor to photoexcite an electron in the photocatalyst material from the valence band to the conduction band, leaving a hole in the valence band. 375,376This photogenerated hole initiates oxidation reactions with the surrounding electrolyte to release reactive redox agents into the solution.Continually regenerated photo-holes and reactive species result in the oxidative destruction of pollutants. 375,377Homogeneous and heterogenous photocatalytic processes have been reported for aqueous AOPs and ARPs.The most often used homogenous process is the photo-Fenton process, which employs iron complexes that undergo photochemical reduction to ferrous iron. 3782][383] Heterogeneous photocatalysis employs solid catalyst materials that are readily recoverable and reusable, reducing separation expenses. 383Halogenated organic pollutants have completely been mineralized to environmentally benign products, using photocatalysis. 376,384,3851][182] Nevertheless, Fenton-based AOPs have extensively been studied for the degradation of brominated and chlorinated pollutants.Fenton-based technology utilizes the reaction of ferrous ions with hydrogen peroxide, at an optimum pH, to generate hydroxyl radicals that react with pollutants in solution. 386While Fenton-based methods are popular due to their wide application range, cost effectiveness of iron, strong anti-interference ability, simple operation, and rapid degradation, 387,388 significant disadvantages exist, such as the narrow working pH range and the generation of significant amounts of iron sludge. 389Brominated flame retardants were completely mineralized, [390][391][392] whereas chlorinated trichloroethylene was not completely degraded. 393Fenton methods in pyrite suspension have been developed to enhance reaction kinetics of H 2 O 2 decomposition to OH radicals for more efficient pollutant degradation. 393Iron ions activated persulfate anions to produce sulfate radicals in a Fenton-like reaction for the chemical degradation of trichloroethylene. 394hotoelectrochemical systems have demonstrated enhanced AOP performance, synergistically degrading halogenated pollutants.Specifically, a photoelectrochemical system has been developed to degrade a series of chlorinated organic molecules to carbon dioxide, carbon monoxide, and chloride ions. 395ulsed potential electrolysis has been shown to significantly increase mineralization efficiencies compared to constant potential electrolysis for chlorinated organic pollutants. 395In a gas diffusion device, 4-chlorophenol and 4-bromophenol were electrochemically dehalogenated to phenol. 175airing Fenton processes with ultraviolet irradiation capitalizes on parallel reactive oxygen generation pathways.An ultraviolet-Fenton system achieved complete degradation of tetrabromobisphenol-A. 396Hybrid Fenton-like processes have been paired with other AOPs and ARPs, such as sonolysis by ultrasound. 397,398Hybrid systems have garnered much attention recently because of enhanced overall degradation rates.Sonophotocatalysis utilizes ultrasonic radiation and photocatalysis to synergistically generate more OH radicals than either individual method by itself does, evidenced by the degradation of 2-chlorophenol. 399Hybrid systems that utilized ozone and H 2 O 2 in a tube reactor nearly completely mineralized trichloroethylene and perchloroethylene in aqueous solution. 400vanced redox processes for fluorinated organic compounds including PFAS Conventional wastewater treatment processes are inefficient for PFAS destruction because of the exceptional stability of C-F bonds (see Table 1 and Fig. 3). 90PFAS can be destroyed by AOPs and ARPs, particularly when aided by photo-assisted processes. 90Direct photolysis is ineffective for treating many PFAS chemicals because the optical absorption of most PFAS molecules is limited to the UV-C region (o220 nm); UV-C irradiation requires specialized equipment and skin-cancer safety precautions; system lifetimes were inherently lower than those of systems irradiated with longer wavelengths, adding to overall capital investments. 90Photocatalysis enhanced PFAS degradation, as was reported for 254 nm (deep ultraviolet) illumination of the benchmark photocatalyst TiO 2 for PFOA decomposition; irradiation with 315-400 nm light significantly decreased degradation performance, suggesting that the ability of the photocatalyst to generate sufficiently strong oxidants, such as OH or OO À , controlled defluorination efficiency. 401][404] Photo-assisted electrochemical processes demonstrated enhanced degradation efficiencies of fluorinated compounds compared to dark electrochemical processes. 260,341,405,406The photo-Fenton process was utilized to increase transient radical concentrations compared to those generated by the dark Fenton process 407 and concomitantly enhanced PFAS degradation. 408ddition of auxiliary chemicals to the electrochemical system benefitted the overall degradation efficiency of PFAS chemicals. 280,330Direct electrochemical and indirect hybrid electrochemical systems have been investigated for fluorinated pollutant degradation. 234,260,280,330,339,341,405,406,409,410Reported direct electrochemical PFAS degradation processes were most efficient when they occurred on boron-doped diamond (BDD) electrodes, 234,339,409,410 which are cost-prohibitive on a large scale (see Fig. 4).

Mechanisms of electrochemical aqueous PFAS destruction
All PFAS electrooxidation pathways start with a direct electron transfer to the anode, which is rate-limiting according to density functional theory calculations, 282 followed by decarboxylation (perfluorinated carboxylic acids, such as PFOA) 176,183,274 or desulfonation (perfluorinated sulfonic acids, such as PFOS) 256,282,341,342 to form a C n F 2n+1 radical, from which CF 2 moieties are unzipped by OH or OO À radicals to form shorterchain perfluoroalkyl radicals, ultimately producing CO 2 and HF, which can safely be mineralized as calcium-containing solids. 411Several reaction pathways have been proposed, depending on the chemical nature of the oxidant (Fig. 7). 304,310,412All reported mechanisms of aqueous electrochemical degradation of PFOA start with electron transfer to the anode and decarboxylation, liberating CO 2 , to form the C 7 F 15 radical, which then further reacts with OH to form C 7 F 15 OH in the hydroxyl radical mechanism or with oxygen and protons in the hydrogen peroxide mechanism. 304,310,412Other auxiliary anions in aqueous solution, such as sulfate or borate, enhance the generation of superoxide radial anions or hydroxyl radicals, opening pathways that proceed through the hydrogen peroxide mechanism or hydroxyl radical mechanism, respectively (Fig. 7A).The intermediate C 7 F 15 OH (structurally equal to C 6 F 13 CF 2 OH) is unstable and undergoes intramolecular rearrangement and hydrolysis, by which C 6 F 13 COO À is produced, effectively unzipping one CF 2 moiety as one CO 2 and two HF molecules from the original C 7 F 15 COO À reactant. 304Alternatively, the C 7 F 15 radical can react with dioxygen in the oxygen mechanism (Fig. 7A) to form the C 7 F 15 OO radical, followed by decomposition of C 7 F 15 O to perfluoorohexyl radical and carbonyl fluoride. 412eported aqueous electrochemical degradation of PFOS starts with a direct electron transfer to the anode.3][414] Nevertheless, mechanism II has been proposed, which unzips CF 2 moieties before desulfonation. 412Desulfonation can follow different mechanistic pathways (Fig. 7B).In mechanism I, the initial electron transfer to the anode is followed by the formation of a very unstable C 8 F 17 SO 3  À radical that reacts with water to form C 8 F 17 , which undergoes hydroxylation and hydrolysis reactions to form deprotonated PFOA and HF.This cycle repeats seven times to produce CO 2 and HF. 413,414 PFOA degradation (Fig. 7A).In mechanism III, PFOS-derived C 7 F 15 has been reported to decompose along the oxygen mechanism pathway of PFOA until CO 2 and HF are obtained as final products (Fig. 7B). 412Likewise, C 7 F 15 decomposition can also proceed through the PFOA hydroxyl radical mechanism to arrive at the desired products CO 2 and HF.Carbon-based radicals are longer-lived than oxygen-based radicals; 418 in aqueous electrolyte, carbon-based radicals are at PFAS molecules, and oxygen-based radicals are the reactive oxygen species that stem from electrolytes.The protonation state, which depends on electrolyte pH and pK a values of reactants and intermediates (see Table 4), critically affects mechanistic pathways.For example, the C 7 F 15 OH intermediate of the hydrogen peroxide and hydroxyl radical mechanisms in Fig. 7A can only be formed at pH values below 2 (see Table 4).

Future research directions
More research is needed to develop viable, globally scalable technologies for the destruction of halogenated organic water pollutants.Systems must work in aqueous media, consist of nonprecious materials, minimize capital expenditures and energy requirements, and be capable of being powered by renewable electricity for sustainability.A quantitative mechanistic understanding of organohalogen destruction via aqueous advanced redox processes is urgently needed, for which analytical detection methods for organohalogens and transient species must be improved, and the development of new catalysts and processes must be advanced.

Analytical PFAS detection methods
Detection of low concentrations of organohalogens in the ppb to ppt range is challenging, particularly in natural water samples, in which ubiquitous abundant ions, such as chloride, can interfere. 419,420Defluorination performance is typically assessed by fluoride ion quantification by ion selective electrode measurements. 421Nuclear magnetic resonance (NMR) spectroscopy has been used since the early 1960s to detect PFOA and PFOS, with a limit of detection of 1.5 mM at that time. 422Gas chromatography (GC) and liquid chromatography (LC) have gradually gained more importance as separation techniques prior to compound identification and quantification.GC can only separate neutral and volatile analytes with low molecular weight, and is limited by low sensitivity and long separation times. 423,424Compared to GC, LC is more widely utilized in quantitative organohalogen analysis, owing to its ability to separate semi-or non-volatile, polar, higher molecular weight, and thermally unstable compounds. 47A widely used analysis technique is LC coupled with mass spectrometry (LC-MS) to detect halogenated contaminants, 425 which is often used together with reference standards that are currently available for one hundred of the hundreds of potentially relevant PFAS that are harmful to human health. 426High-resolution mass spectrometry (HRMS) has emerged as a key tool for identifying legacy and novel PFAS.HRMS can detect more than 750 PFAS. 427LC and GC, coupled with various HRMS-based techniques, have been widely used for the identification and quantification of organohalogens. 428LC/GC-HRMS has extremely high selectivity, high resolving power, and is capable of detecting unknown PFAS.The lowest limit of PFAS detection for LC/GC-HRMS that has been reported was 100 nM. 429The high cost of HRMS instrumentation and challenges with respect to differentiation of PFAS isomers are the main disadvantages of LC/GC-HRMS. 26PFAS concentrations down to 10 nM are detectable by mobility spectrometry combined with LC and MS, which provides high selectivity and sensitivity as well as fast detection, but is expensive and unable to distinguish PFAS isomers, which limits broad usability. 430

Catalyst development
The development of new, nonprecious, efficient, selective, and robust catalysts is critical for the advancement of halogenated pollutant destruction.Catalyst materials are the foundation of degradation technologies because catalysts lower kinetic barriers in the conversion of supplied energy into chemical energy that initiates the degradation process.6][437][438][439][440][441] Understanding how surface species bind to catalysts through first principles and machine learning computations is key to inform and direct experimental investigations in the vast materials space. 442,443he electronic and geometric structure of catalysts, including changes during turnover and bonding motifs of reaction intermediates can be determined by advanced in situ or operando spectroscopies, combined with first principles and machine learning.In situ techniques monitor changes of the catalyst material under turnover, whereas operando spectroscopies additionally identify and quantify generated products during catalysis, thus enabling direct catalyst property-performance relationships during catalytic cycling.In situ and operando spectroscopies are often employed to gain a mechanistic understanding of electron, ion, and mass transport at catalyst interfaces, 444 providing insights into the surface and bulk structure of catalysts, their composition, oxidation states, and adsorbed intermediates under reaction conditions. 436,441,445 situ and operando spectroscopies additionally enable critical insights into reaction pathways, 446 especially when coupled with computational studies.Some of the most frequently used in situ and operando characterization techniques of catalyst materials are in situ Raman and surface enhanced Raman spectroscopy (SERS), 445,447 Fourier transform infrared spectroscopy (FTIR), 446 and the synchrotronbased techniques X-ray diffraction, 446 X-ray absorption spectroscopy, 448 near ambient pressure X-ray photoelectron spectroscopy, 446 high-energy-resolution fluorescence-detection X-ray absorption spectroscopy, 449 and X-ray absorption fine structure spectroscopy, 449 sometimes used in grazing incidence angle configuration to enhance catalyst surface specific information. 450o gain a more accurate and quantitative understanding of catalysts and to reveal critical insights into the thermodynamics and kinetics of species during electrocatalysis, theoretical approaches, such as first principles density functional theory (DFT) calculations and microkinetic modelling simulations are often combined with in situ and operando characterization techniques. 436,444,451,452First-principles calculations, rooted in quantum mechanical electronic structure theory, have been utilized to analyse the intimate and dynamic relation between the microscopic processes and the meso-to macroscopic environment. 453DFT is the most commonly used first-principle method and utilizes functionals of spatially dependent electron density calculations to investigate the electronic structure of atoms, molecules, and solids. 454First-principle calculations are especially important in the context of catalyst discovery because they enable screening for new catalysts based on user-defined criteria and hypotheses, such as optimal activity, stability of elements under specific conditions, formation energy, selectivity, and material phases. 2,455These computational screening tools have been employed to analyze ensemble effects and electronic effects, to determine catalytic activity and selectivity. 455Electronic effects control the binding of the reaction intermediates. 455DFT calculations can be used to predictively design advanced catalysts, utilizing functional mechanisms of existing catalysts, comparing them, and making predictions which of these mechanisms is most suitable for each future reaction. 442,455Theoretical predictions from first-principle calculations are based on properties calculated from basic physical quantities and do not consider experimental results.Thus, it is challenging to quantitatively understand catalysts by advanced in situ or operando spectroscopies or theoretical models alone.Therefore, experimental results and theoretical models are nowadays often combined with machine learning algorithms that enable a better understanding of the nature of chemical bonding and its variation in strength across physically tuneable factors. 442Data-driven artificial intelligence models are capable of being integrated into active and iterative learning schemes that incorporate experimental results to improve the extrapolative, i.e., predictive, capabilities of models. 456,457Machine learning can also be combined with multiscale simulations and quantum mechanics to predict the performance of surface sites of catalysts. 458This journal is © The Royal Society of Chemistry 2023 approaches enable much-needed quantitative understanding of catalytically active sites, reactions centres, and reaction mechanisms. 446,451,458,459Right now, detailed DFT calculations and machine learning studies have challenges, such as high computational cost and potential loss of physical intuition since most models consist of complicated mathematical formulations that are difficult to interpret. 442Furthermore, reporting of experimental and computational information in much detail must be ensured to overcome challenges regarding choosing suitable functionals in DFT calculations. 460,461ulti-disciplinary approaches that combine experimental electrochemical performance assessments, operando spectroscopies, and computational approaches have to date the greatest potential to advance catalyst development. 451

Process development
The thermodynamics, kinetics, mass transport, and chemistries of a system (i.e. the four pillars of chemical engineering) govern the overall electrocatalysis process and are tunable by the electrocatalytic process parameters outlined in Fig. 8.The electrolyte pH critically affects the thermodynamic potentials and chemistries, including generation mechanisms and protonation state, of reactive species produced during catalytic turnover (see Table 3).Electrolytes composed of small ions enhance mass transport, and electrolyte composition affects chemistries, particularly with respect to auxiliary solutes that can form transient radicals with high oxidation or reduction strengths.The electrolyte temperature impacts kinetics and mass transport of the system.The reaction time influences reaction kinetics, and electrolyte agitation increases mass transport.The applied potential controls the thermodynamics, kinetics, and chemistries by making different mechanistic pathways accessible.
The scaleup of halogenated pollutant destruction technologies from the laboratory scale to industrially relevant scales demands that investigated systems are operated in water at pollutant concentrations that occur in the field, or at least at concentrations that are achievable by separation techniques, 462,463 necessitating optimization of substrate mass transport.

Detection of transient ARP and AOP species
Quantitative analytical techniques are needed for the detection of short-lived, highly reactive redox species 352,[464][465][466] that are produced during assisted or unassisted electrocatalytic ARPs and AOPs, to develop a quantitative mechanistic understanding of organohalogen destruction processes.8][469] Freeze or spin trapping preserves radical species long enough to detect them in an EPR spectrometer. 335Spin traps are chemicals that react with radicals to create longer-lived, paramagnetic, EPR-active species, which are used to identify radicals with short lifetimes. 335ualitative radical identification is important for understanding which species are involved in ARP and AOP mechanisms.Nevertheless, quantification is necessary to deeply understand mechanisms, and optimize reaction conditions and overall pollutant degradation.While EPR is typically used for identification of radical species, an external standard, such as 2,2,6,6tetramethylpiperidine 1-oxyl (TEMPO), can be utilized to obtain quantitative data. 470Other quantitative detection methods are fluorescence and optical spectroscopy that utilize turn-on or turn-off dyes to determine radical concentrations. 352,471,472hese dyes typically react directly with radical intermediates, resulting in a change in the fluorescence or absorption spectrum of that dye, enabling the determination of concentrations of radical intermediates.Faster, more cost-effective, and more sensitive detection methods must be developed to accelerate the innovation of viable aqueous electrocatalytic organohalogen destruction techniques.

Conclusions
Halogenated organic compounds are widespread due to their exceptional utility and because they were initially considered nonmetabolizable and nontoxic, owing to the extreme stability of carbon-halogen bonds.However, prolonged use has led to their accumulation in the environment and organisms globally.We have introduced various classes of these pollutants categorized by halogen type (fluorine, chlorine, bromine), discussed important policies and regulations, outlined their applications, and highlighted associated environmental and human health risks.We discussed remediation techniques, focusing on carbon-halogen bond strengths, capital expense and energy needs for destruction, and electrocatalytic aqueous advanced redox processes.We highlighted mechanistic details of electrocatalysis, including oxidations and reductions of the water-oxygen system, as well as thermodynamic potentials, protonation states, and lifetimes of radicals and reactive oxygen species in aqueous electrolytes, importantly, at different pH conditions.We pointed out that advanced reduction processes necessitate anaerobic conditions, which are impractical beyond the laboratory scale because electrocatalytic water oxidation can produce dioxygen, membranes partitioning oxidation and reduction half-reactions cannot completely prevent oxygen crossover, and airtight seals are inherently challenging in large-scale applications; ergo, advanced oxidation processes appear to be more promising.We reviewed aqueous advanced redox processes for different halogenated compounds and PFAS (per-and polyfluoroalkyl substances), detailing potential mechanisms.Future research directions require quantitative understanding of destruction mechanisms, improved detection methods, advanced catalyst development, and energy-efficient processes.Scalable systems using nonprecious materials, powered by renewable electricity, are crucial.We outlined the interconnectedness of electrocatalytic process parameters and their effect on the chemical engineering descriptors thermodynamics, kinetics, mass transport, and chemistries, all vital in steering complex reaction networks.Finally, we suggested strategies to accelerate the development of effective aqueous electrocatalytic techniques for organohalogen destruction.
, and came to the U.S. for her undergraduate education at the University of Rochester, where she obtained dual-major BS and BA degrees in chemical engineering and dance in 2023.At the University of Rochester, she conducted undergraduate research in the group of Prof. Astrid M. Mu ¨ller, working on aqueous electrocatalytic defluorination of the PFAS chemical perfluorooctanoic acid, catalyzed by nanomaterials made by pulsed laser in liquids synthesis.In fall 2022, Wanqing received an Undergraduate Teaching Assistant Award in chemical engineering.Currently, Wanqing is a PhD student in chemical engineering at Georgia Institute of Technology.Astrid M. Mu ¨ller Astrid M. Mu ¨ller is an Assistant Professor of Chemical Engineering at the University of Rochester since 2018.Prof. Mu ¨ller earned a PhD in Physical Chemistry for work on ultrafast reaction dynamics at the Max Planck Institute of Quantum Optics.Her postdoctoral work centred on developing a fundamental understanding of laser-matter interactions.Her independent research focuses on pulsed laser in liquids synthesis of mixed-metal nanomaterials with controlled structural and electronic properties.This uniquely positions Prof. Mu ¨ller's group to quantitatively understand how nanocatalysts and electrocatalytic mechanisms impact the performance of nanomaterials in sustainable energy, green chemistry, and aqueous PFAS destruction applications.Ziyi Meng Ziyi Meng obtained his BSc in Polymer Materials and Engineering (2019) and MSc in Materials Science (2022) at Hubei University of Technology (China), where he conducted research in microbial modified mineral materials, advanced functional super elastic, and inorganicorganic composited proton exchange membrane fuel cells for extreme conditions under the supervision of Prof. Qingting Liu.In 2022, he enrolled at the University of Rochester as a PhD student in the Materials Science Program and joined the group of Prof. Astrid M. Mu ¨ller to work on the degradation of per-and polyfluoroalkyl substances (PFAS), employing aqueous electrocatalysis and laser-made bimetallic nanomaterials.

Fig. 2
Fig.2Human health risks upon exposure to halogenated organic water pollutants.

Fig. 6
Fig. 6 Generation of radicals and reactive species in aqueous electrocatalysis for use in advanced redox processes; ref. el., reference electrode.Species produced at electrodes are in white, species produced in bulk solution are in blue in the blue circles; the yellow stars indicate assisting processes, such as deep ultraviolet light irradiation, radiolysis, or sonolysis.Anodic dioxygen (O 2 ) may be formed, depending on the water oxidation catalyst.Species in braces are short-lived intermediates.(A) Electrocatalysis in aerobic electrolyte, (B) electrocatalysis in anaerobic electrolyte, (C) assisted electrocatalysis in aerobic electrolyte, (D) assisted electrocatalysis in anaerobic electrolyte; e À aq , hydrated electron.

Fig. 8
Fig. 8 Electrocatalytic process parameters, color-coded to visualize which chemical engineering descriptor they affect.

Table 1
Bond dissociation energies for different binding motifs of halogens X for common organohalogens.Energy ranges originate from different bond dissociation energies depending on the position of the F atom within the molecule.PFOA, perfluorooctanoic acid; PFOS, perfluorooctane sulfonic acid.From ref.116-123

Table 3
Reduction reactions of the O 2 -H 2 O-system and thermodynamic potentials E at pH 0, 7, and 14.From ref.309, 315 and 316 aq This journal is © The Royal Society of Chemistry 2023 Assisted aerobic electrocatalysis creates OH radicals, H 2 O 2 , or O 2 anodically by water oxidation, and H 2 O 2 , OH, or OO À by enable assisted electrocatalytic production of reactive oxygen radicals and, in O 2 -free aqueous environment, hydrated electrons, e À aq , and hydrogen radicals, H .
In the presence of hydrogen radicals (H ), aqueous electrocatalytic PFOS degradation follows mechanisms II or III, where PFOS (C 8 F 17SO 3 417 whoseThis journal is © The Royal Society of Chemistry 2023 spectroscopic signature is known,417and the cycle repeats seven times until only CO 2 , HF, and SO 3 À remain of PFOS.In contrast, in mechanism III, the intermediate C 8 F 16 SO 3 À undergoes an H/F exchange and a decarboxylation reaction to form C 7 F 15 , which is an intermediate of aqueous electrocatalytic